Journal of Fish Biology (2003) 62, 1392–1404
doi:10.1046/j.1095-8649.2003.00124.x, available online at http://www.blackwell-synergy.com
Effects of gizzard shad on benthic communities in
reservoirs
K. B. G I D O
University of Oklahoma, Biological Station and Department of Zoology Norman,
OK 73019, U.S.A.
(Received 20 November 2002, Accepted 14 April 2003)
Effects of gizzard shad Dorosoma cepedianum on benthic communities in a large southern
reservoir (Lake Texoma, U.S.A.) were examined during two field enclosure and exclosure
experiments in which enclosures were stocked at high and low densities in 1998 and 1999,
respectively. In both years, chironomid abundance significantly increased in treatments that
excluded large fishes from foraging on sediments. Mean abundance of chironomids and
ostracods were significantly higher (P < 005) in exclosures than enclosures stocked with gizzard
shad at 1140–1210 kg ha1. In 1999, benthic invertebrate abundances did not differ (P > 008)
between exclosure and enclosures stocked at 175–213 kg ha1. Per cent organic matter, algal
abundance and abundance of other macroinvertebrates in sediments did not differ significantly
among treatments in either year. Although chironomid abundance was reduced in gizzard shad
enclosures in 1998, food habits from this and other studies showed that adult gizzard shad in
Lake Texoma only consumed detritus and algae. It is likely that high sedimentation rates in
Lake Texoma limit the ability of gizzard shad to regulate algae and detritus in benthic
sediments. Thus, it is concluded that disturbance of benthic sediments by gizzard shad caused
the observed reduction in chironomid abundance, rather than through consumption or compe# 2003 The Fisheries Society of the British Isles
tition for resources.
Key words: algae; benthic organisms; chironomids; detritus; omnivorous fishes; reservoirs.
INTRODUCTION
The ability of benthic fishes to regulate ecosystem processes in reservoirs has
important management implications because these fishes may indirectly influence
other economically important species (e.g. sport fishes) by altering the prey base or
productivity in those systems. For example, benthic foraging by large-bodied
omnivorous fishes stirs up benthic sediments, increases water-column nutrient
concentrations and increases rates of nutrient regeneration (Lamarra, 1975;
Shormann & Conter, 1997). Moreover, planktivory and benthic foraging by gizzard shad Dorosoma cepedianum (Lesueur) allows them to regulate zooplankton
abundance through both top-down and bottom-up processes (Vanni, 1996;
Schaus & Vanni, 2000). Because the general effects of benthic fishes are thought
Present address: Kansas State University, Division of Biology, Manhattan, KS 66506, U.S.A.
Tel.: þ1 785 532 5088; fax: þ1 785 532 6653; email: kgido@ksu.edu
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FISH EFFECTS ON BENTHIC COMMUNITIES
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to negatively influence predatory sport fishes, a common management objective is
to eliminate or reduce their abundance (Wydoski & Wiley, 1999; Tomlinson et al.,
2002).
In addition to changes in water chemistry and nutrient dynamics, benthic fishes
can also reduce the abundance of benthic invertebrates. In general, most studies
in lentic systems that have excluded fishes from foraging on sediments reported
an increase in invertebrate abundance in the absence of fishes (Gerking, 1994).
One exception was Thorp & Bergey (1981) who concluded that environmental
heterogeneity and food web complexity limit the ability of vertebrates (fishes and
turtles) to control benthic invertebrate populations, and that other factors such as
temperature and habitat were more important regulators of these organisms.
Several other studies using field enclosures or exclosures have reported minimal
effects of fish on benthic invertebrates in streams with coarse substrata (Allan,
1982; Reice & Edwards, 1986) and ponds with high densities of macrophytes
(Gilinsky, 1984; Batzer et al., 2000). Field cage experiments, however, have
indicated a strong effect of fishes on benthic invertebrate communities in farm
ponds (Morin, 1984) and warm-water, soft-sediment streams (Gilliam et al.,
1989). Moreover, the exclusion of fishes in streams has been shown to directly
or indirectly affect invertebrates, algae and detritus (Power et al., 1985; Flecker,
1996). Indeed, the relative importance of fishes varies across habitats and it is
necessary to perform experiments across a broad range of systems and species
groups to predict the conditions under which fishes are important regulators of
benthic community composition.
The ability of a species to regulate properties of an ecosystem may depend on
its abundance (Power et al., 1996; Power, 1997), population size structure
(Mehner et al., 1998), trophic level (Carpenter et al., 1992; Schindler et al.,
1993) or mode of feeding (Matthews, 1998). Large-bodied benthic fishes that
can process large amounts of sediments (e.g. Prochilodus, Flecker, 1996) are
likely to disrupt sediment processes more than disturbances by smaller fishes.
North American reservoirs often have high densities of large-bodied benthic
omnivorous fishes, including gizzard shad. Because adult gizzard shad are large
enough to escape predation by most predators, their populations have expanded
and they are typically the numerically dominant fish species in many North
American reservoirs (Stein et al., 1995).
Recent studies indicate gizzard shad can potentially have strong effects on
ecosystem functioning of reservoirs. Gizzard shad are facultative detritivores,
but primarily consume detritus and algae as adults (Dalquest & Peters, 1966;
Gido, 2001). Mundahl (1991) estimated that dry mass of sediments processed
daily by gizzard shad was equivalent to 13% of their wet body mass. Because of
high sedimentation rates, however, he suggested gizzard shad were unlikely to
affect benthic community dynamics. By foraging on sediments, gizzard shad
also transport nutrients into the water column of reservoirs through excretion;
this flux of nutrients can contribute substantially to the total nutrient budget of
the reservoir during periods of low inflow from tributaries (Vanni, 1996; Schaus
et al., 1997; Schaus & Vanni, 2000; Gido, 2002). As far as is known, only one
study has examined the effects of gizzard shad foraging on benthic invertebrate
abundance. Using 900 l mesocosms stocked with gizzard shad at c. 1000 kg ha1,
Cline et al. (1994) found no significant effect of gizzard shad on the density of
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K. B. GIDO
benthic invertebrates. Because these were closed systems, however, indirect
effects of the fish (e.g. increased nutrient levels) may have obscured the effects
on the benthos.
In this study, field enclosure and exclosure experiments were conducted to
evaluate the effects of gizzard shad on benthic processes in Lake Texoma
(33 490 N; 96 330 E), a large impoundment on the Oklahoma–Texas border,
U.S.A. Gizzard shad are abundant in this system and comprised c. 14% of
the offshore fish biomass captured during gillnet surveys (Gido & Matthews,
2000; Gido et al., 2000). Thus, gizzard shad are probably an important component of the benthic food web that, based on their feeding ecology, are
hypothesized to either directly or indirectly affect benthic invertebrates, algae
and detritus. Direct effects should occur through consumption or mechanical
disturbance of the sediments and indirect effects would probably occur through
exploitative competition. It is predicted that the exclusion of gizzard shad from
feeding on sediments will cause an increase in algae, detritus and benthic
invertebrate abundance.
MATERIALS AND METHODS
STUDY AREA
Lake Texoma is a 36 000 ha impoundment of the Washita and Red Rivers. Reservoir
releases and resulting fluctuations in water level are primarily for hydropower and flood
control. Secchi disk depth transparency typically ranges from 100 to 125 cm, but can
decrease to 15 cm during turbid inflow episodes (Matthews, 1984). Locations of fish
collections and the field experiment were c. 35 km upstream from Denison Dam, within
the Red River arm of Lake Texoma. Experimental cages were placed in a shallow cove
(Mayfield Flats) west of the University of Oklahoma Biological Station that was sheltered from the prevailing south winds (Patten, 1975). This silt- and sand-bottomed cove is
typical of other coves in the transitional zone of the reservoir (Kimmel & Groeger, 1986).
Catches from benthic-set gillnets in this cove during summer and autumn of 1997 and
1998 (Gido & Matthews, 2000) showed that gizzard shad numerically dominated the
assemblage (44%). Striped bass Morone saxatilis (Walbaum 16%), threadfin shad Dorosoma petenense (Günther 14%) and smallmouth buffalo Ictiobus bubalus (Rafinesque)
(6%) were commonly collected.
EXPERIMENTAL DESIGN
Experimental cages were constructed with 254 cm mesh plastic netting to allow rapid
exchange of water through the cage walls. Although cages were sheltered from direct
exposure to wind-waves, there were sufficient currents (i.e. ‘fetch’ within the cove) to
inhibit the accumulation of excreted nutrients and potentially faeces within the cages.
This mesh-size also allowed smaller fishes [e.g. juvenile Lepomis and Morone spp.,
Menidia beryllina (Cope) and Pimephales vigilax (Baird & Girard)] to pass through
while retaining or excluding the large-bodied (>100 mm total length, LT) focal species.
Cages were open at the top and bottom and edges were secured to the substratum with
metal posts and rocks. All cages were carefully placed in c. 15 m water with minimal
disturbance to the enclosed sediments. Fish were captured with a 40 m 15 m beach
seine and immediately placed in cages. Two separate experiments were conducted during
July 1998 and May 1999. In the first experiment, the effects of gizzard shad were tested at
relatively high densities (1140–1210 kg ha1). For this experiment, eight cylindrical cages,
each 194 m in diameter, were used to either enclose or exclude fish from foraging on
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benthic sediments. The experiment was a factorial design with three treatments and four
replicates each. As a measure of ambient conditions, one treatment included samples
taken in a haphazardly chosen location in close proximity (<15 m) to the cages. By
taking these ambient samples close to cages, this controlled for a potential effect of cage
walls. In a second exclosure treatment no fish were added. The third treatment enclosed
three gizzard shad (211–241 mm standard length, LS).
Cove rotenone surveys of North American reservoirs by the National Reservoir
Research Program (unpubl. data) estimated that the mean S.D. standing crop of gizzard
shad was 119 176 kg ha1 and ranged from 0 to 2232 kg ha1 for 360 reservoirs.
Average standing crop estimates for gizzard shad in Oklahoma reservoirs based on
cove rotenone surveys was 118 kg ha1 (Jenkins, 1976). Thus, densities used in the
above experiment probably represent an upper limit for Lake Texoma and a second
experiment was conducted in 1999, using larger cages (388 m diameter) stocked with four
individuals per cage (225–240 mm LS, 175–213 kg ha1). The 1999 experiment was the
same factorial design with three treatments (fish enclosure, ambient and fish exclosure),
but had three replicates per treatment. At the end of each experiment, a 3 m 15 m seine
with 5 mm bar mesh was used to collect all fishes (including those fishes that could
passively move into cages). The entire intestinal tract of gizzard shad was dissected and
per cent composition of diet by volume was estimated in the laboratory under 40
magnification following methods of Gido (2001).
R E S PO N S E V A R I A B L E S
To examine changes in sediment biota, core samples (8 cm diameter) were taken from
a haphazardly selected position within each cage and at locations outside of cages at
the beginning of the experiment and every 4 days thereafter. To minimize the disturbance
of substrata, only one core sample was taken per sampling interval in the 194 m diameter
cages in 1998 and three cores (pooled for each cage) were taken in the 388 m diameter
cages in 1999. A previous study (Gido, 2001) indicated invertebrate abundance in
replicate sediment cores are very homogeneous throughout this and two other coves
in this region of the reservoir (CV of replicate samples averaged 21%), probably because
of the even distribution of fine sediments in these sheltered coves. Thus, this appeared to
be an adequate sample to characterize the benthic communities in these cages. Only
the upper 10 mm of core sediments were retained because organisms and detritus
deeper in sediments are most likely inaccessible to these fishes (Mundahl & Wissing,
1987). Samples were preserved in 5% formalin. In the laboratory, water was added to
each sample to bring it to a total volume of 200 ml. This sample was shaken vigorously
to suspend all silt, sand and organic matter, and four, 4 ml aliquots were drawn from
this homogenate. The remainder of the sample was passed through a 210 mm sieve to
retain macroinvertebrates and large zooplankton. One of the 4 ml aliquots was placed in
an aluminium tray, dried at 60 C for 24 h and then ashed at 550 C for 1 h to determine
per cent organic matter, a composite for algae and detritus. Algal abundance was
determined from the other three aliquots (i.e. three replicates per core sample). A
01 ml sub-sample from each aliquot was placed in a Palmer counting cell and algal
cells were counted in 50 fields of view at 400 magnification (Flecker, 1996). Filamentous algal strands were considered one cell because they were small and varied little
in size.
DATA ANALYSIS
Differences in response variables among treatments in the experiments were assessed
with a repeated measures ANOVA with time as the repeated factor. Because core samples
were taken prior to the addition of fish on day 0, this date was excluded from the
analysis, but the data are plotted for reference. If significant differences were found
among treatments, post-hoc comparisons were made with Ryan’s multiple comparison
procedure (Toothaker, 1991).
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RESULTS
Core samples were overall, numerically dominated by several genera of
chironomids (primarily Dicrotendipes spp. and Glyptotendipes spp.), which
accounted for 67% of the individuals averaged across years and treatments.
Ostracods (16% of total individuals) and oligochaetes (15% of individuals) were
commonly found. Less abundant taxa in core samples included copepods, water
mites (Hydracarina) and phantom midges (Chaoborus spp.). These less abundant taxa were excluded from the analysis. Algal counts were dominated by
diatoms with few filamentous green algae and cyanobacteria (67, 19 and 14% of
total biovolume averaged across years and treatments).
In 1998, mean chironomid densities differed significantly among treatments
(F2,8 ¼ 790, P ¼ 0013; Fig. 1). Because the effects of time and the interaction
between time and treatment were not significant, post-hoc comparisons were
only made among the three treatments. In this analysis, densities of chironomids
in ambient and enclosure treatments were lower than in core samples from
FIG. 1. Mean þ S.E. chironomid densities in benthic core samples taken from experimental treatments in
(a) 1998 and (b) 1999. &, Ambient; &, exclosure; , Dorosoma cepedianum.
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exclosures; there was no difference between ambient and enclosure treatments.
Ostracod density differed across time (F4,32 ¼ 2970, P ¼ 0034) and among
treatments (F2,8 ¼ 1311, P ¼ 0019), but the interaction of these variables was
not significant (Fig. 2). As with chironomid densities, post-hoc comparisons
among treatments indicated that ostracod densities were greater in the exclosure
than ambient or enclosure treatments, but not different between ambient and
enclosures. Although there was a significant effect of time on oligochaete
density (F4,32 ¼ 377, P ¼ 0013), there was no treatment or time treatment
interaction (Fig. 3). Algal biovolume (not shown) and per cent organic matter
(Fig. 4) did not differ across time or among treatments (all P-values > 01).
Chironomid densities in the 1999 experiment were significantly different
among treatments (F2,6 ¼ 339, P < 0001; Fig. 1). The effect of time and the
time treatment interaction were not significant. Post-hoc comparisons among
treatments revealed that chironomid abundance was significantly lower in
ambient treatments than gizzard shad enclosures and exclosures. Although
FIG. 2. Mean þ S.E. ostracod densities in benthic core samples taken from experimental treatments in
(a) 1998 and (b) 1999. &, Ambient; &, exclosure; , Dorosoma cepedianum.
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K. B. GIDO
FIG. 3. Mean þ S.E. oligochaete densities in benthic core samples taken from experimental treatments in
(a) 1998 and (b) 1999. &, Ambient; &, exclosure; , Dorosoma cepedianum.
mean chironomid densities in enclosures were c. 60% lower in enclosures
than exclosures on day 16 and 30% lower on day 20; these differences were
not significant (P ¼ 0081), possibly because of the low power of this experiment
(1 – b ¼ 0237). Ostracod density did not differ across time or among treatments
(all P-values > 01; Fig. 2) nor did algal biovolume (all P-values > 025). Oligochaeta density (F4,24 ¼ 914, P < 0001; Fig. 3) and per cent organic matter
(F4,24 ¼ 522, P ¼ 0009; Fig. 4) both showed a significant decrease across sample
dates but no treatment effect.
Food habits of gizzard shad taken from enclosures at the end of the experiment consisted primarily of detritus and algae (982% by volume) with a few
ostracods (16%) and no chironomids. Gizzard shad appeared to consume
primarily vegetative detritus and diatoms rather than filamentous algae.
The presence of small fishes that could pass through the cage mesh could
have affected the abundance of benthic organisms and organic content of
sediments inside cages. Although there was no significant difference between
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FIG. 4. Mean þ S.E. per cent organic matter in benthic core samples taken from experimental treatments
in (a) 1998 and (b) 1999. &, Ambient; &, exclosure; , Dorosoma cepedianum.
enclosures and exclosures or between years in mean total density of smaller
fishes (ANOVA, P < 03), there were relatively high densities in all cages; bluegill
Lepomis macrochirus Rafinesque, Morone spp. and bullhead minnow P. vigilax
mean S.D. densities were 144 95, 44 31 and 11 06 individuals m2,
respectively.
DISCUSSION
Chironomid abundance was strongly affected by experimental treatments in
this study, but these effects varied between the two years. In both 1998 and 1999
experiments, ambient chironomid densities were approximately three times
lower than in fish exclosures by day 12, probably because of the exclusion of
large-bodied fishes that prey on chironomids or disturb sediments. In addition,
ostracods were shown to increase in exclosures in the 1998 experiment, however,
this effect was strongest on day 8 and was absent by day 20. These differences
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K. B. GIDO
suggest that either a single species or a combination of species are affecting
benthic invertebrate abundance. Moreover, data from the fish enclosures suggest that gizzard shad are at least partially responsible for this pattern, but there
are clearly other large-bodied species, e.g. channel catfish Ictalurus punctatus
(Rafinesque) and smallmouth buffalo in the system that, when excluded from
the sediments, allowed chironomid densities to increase.
Treatment effects in this study may have been obscured because of temporal
differences in chironomid abundance and sediment organic content between
years; chironomid densities were approximately twice as high and organic
content was more than three times higher in May 1999 than in July 1998.
Differences in chironomid densities probably represent a seasonal pattern, as
a decline in chironomid abundance from May to August is typical in Lake
Texoma (Gido, 2001). Rising water levels in 1999 and dropping water levels in
1998 may have been partly responsible for differences in organic content of
sediments between years. Although it is unclear what effect these factors had on
the experimental results, relative difference between exclosures and ambient
conditions were similar between years. Thus, the interannual differences were
probably minimal.
Results from this study contrast with those by Cline et al. (1994) who found no
difference in benthic invertebrate densities between control and gizzard shadstocked mesocosms. Because the current study was conducted in larger field
cages, it is possible that processes differed among these experiments. For example,
nutrients or faeces released by gizzard shad, which could have indirectly increased
benthic invertebrate abundance, might be expected to positively affect benthic
invertebrate abundance in mesocosms. In open field cages, however, any effect of
nutrient or faecal accumulation would be diluted by high exchange rates with
surrounding water.
A number of other studies in which the abundance of benthic fishes in lentic
environments has been manipulated also reported increased densities of benthic
invertebrates in the absence of fishes (Gerking, 1994). Perhaps this is a general
pattern for reservoirs as well (Wetzel, 2001). For example, removal of Catostomus
commersoni (Lacépède) from a Michigan lake was associated with a 13–18-fold
increase in abundance of chironomid larvae (Hayes et al., 1992). Moreover,
a fish kill of Abramis brama (L.) in Lake Ringsjon resulted in an increase in the
benthic invertebrate fauna of this lake (Bergman et al., 1999). The strong
negative effect of benthic fishes on sediment-dwelling invertebrates is probably
because of several processes including direct predation and mechanical disturbance of substrata.
Examination of gut contents at the end of the experiment, in addition to
other published reports, suggests that adult gizzard shad primarily forage on
detritus and algae (Dalquest & Peters, 1966; Mundahl & Wissing, 1987; Gido,
2001). Thus, the reduction in chironomid abundance in gizzard shad enclosures
did not appear to be caused by direct predation. Because the dominant species
of chironomids in Lake Texoma are collector-gatherers that feed on detritus
and algae (Vaughn, 1982), it is possible that gizzard shad may compete for
resources (e.g. diatoms and detritus) with chironomid larvae. This would require
that resources are limited, however, which does not appear to be the case in
Lake Texoma as organic matter and algal biovolume (volumetric quantity of
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algae in or attached to benthic sediments) did not differ among treatments in
either year. Because there was no apparent reduction of detritus or algae in
enclosures, resources did not appear to be limited and competition for resources
does not appear to explain the observed effect. The lack of decline in algae or
organic matter in the gizzard shad enclosures relative to exclosures was surprising because this species foraged almost exclusively on these items. Several
factors could have contributed to a weak association between the presence of
fishes and abundance of benthic algae and organic matter. Foremost, is that
high sedimentation rates of silt, detritus and algae may have dampened the
effects of fishes. For example, Mundahl (1991) predicted that gizzard shad
processed <4% of the deposited sediments in Acton Lake, Ohio. Typical of
many reservoirs, Lake Texoma is very turbid and sedimentation rates were
estimated as between 150 and 550 g dry mass m2 day1 in the region of the
reservoir where this study was conducted (Gido, 2001). The continuous ‘rain’
of organic matter from the water column would probably buffer the removal of
these materials by fishes. Thus, it can be hypothesized that reduced densities of
chironomids in gizzard shad enclosures was caused by mechanical disturbance
of substrata.
Whereas foraging by stream fishes is known to reduce structural habitat
(e.g. filamentous algae and detritus) for invertebrates (Gelwick & Matthews,
1992; Flecker, 1996), foraging by benthic fishes in soft sediments did not appear
to influence the physical structure of the sediments. The mechanical mixing of
the sediments, however, probably alters the chemical exchange equilibria at
the sediment–water interface and changes the microhabitat for these small
benthic organisms. Several studies of marine fishes and spawning salmonids
have reported the effects of foraging-related disturbances on benthos (Virnstein,
1977; McCormick, 1995; Peterson & Foote, 2000). Palmer (1988) found
that disturbance by predatory marine fishes accounted for virtually all the
nematode mortality in laboratory experiments. Experiments in freshwater
lentic systems report strong effects of common carp Cyprinus carpio (L.)
and small mouth buffalo on resuspension of sediments and changes in nitrification rates (Richardson et al., 1990; Shormann & Conter, 1997). In addition,
the stirring of sediments during feeding may reduce chironomid abundance
by dislodging individuals which are then either swept away by currents or
consumed by other fishes. Although direct experimental evidence is lacking
to show that mechanical disturbance of sediments was responsible for the
reduced densities of chironomids in fish enclosures relative to exclosures, this
appears to be the most parsimonious explanation for the observed treatment
effects.
Based on this and other studies (Andersson et al., 1978; Hayes et al., 1992;
Bergman et al., 1999), benthic fishes appear to have a capacity to influence
benthic invertebrate composition in soft-bottomed littoral habitats of reservoirs.
The effect of these fishes, however, is highly dependent on density. In this study
chironomid and ostracod densities in gizzard shad enclosures stocked at
175–213 kg ha1 were not significantly different than exclosures. Although the
statistical power of this comparison was low and the treatment effect was only
marginally insignificant, gizzard shad appeared to have a weaker effect on the
benthos at lower densities. Assuming that these results apply to other reservoirs
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and that standing crop estimates for North American reservoirs averages
119 kg ha1 (National Reservoir Research Program, unpubl. data), it would
appear that most reservoirs do not have sufficient gizzard shad densities to
significantly reduce chironomid abundance. There is, however, much variation
in abundance across years and longitudinally within reservoirs and, as reported
here, gizzard shad can have rapid effects (<20 days) on the benthos. Thus, it is
likely that benthic invertebrates in reservoirs that experience high gizzard shad
densities, even temporarily, may be limited by their feeding activities.
Benthic invertebrates can be affected by top-down processes such as predation and bioturbation (disturbance of sediment layers by biological activity) as
well as bottom-up processes such as sedimentation of algae and detritus. In
Lake Texoma, high sedimentation rates provide adequate resources to allow
benthic invertebrates to increase when fishes are excluded from feeding on
sediments. Thus, reduction or removal of gizzard shad in this and other reservoirs would probably result in an increase in abundance of benthic invertebrates, at least over short time scales. An increase in the biomass of benthic
invertebrates resulting from the removal of benthic fishes, however, may not be
transmitted to higher trophic levels (Hayes et al., 1992). Future studies should
consider the net effect of these fishes in the ecosystem based on a variety of
pathways (e.g. trophic and nutrient dynamics) before fish removal is considered
as a means to increase sportfish production.
This would not have been possible without field and laboratory assistance from
D. Cobb, A. Gido, R. Gido, A. Marsh, D. Lutterschmit and J. Schaefer. I extend special
thanks to W. Matthews for thoughtful discussions and his critical review of this manuscript. This manuscript further benefited by critical reviews from L. Canter, C. Hargrave,
M. Kaspari, N. Mundahl, J. Schaefer, W. Shelton, D. Spooner, C. Vaughn, M. Weiser
and two anonymous reviewers. The University of Oklahoma Biological Station provided
logistic and financial support. Thoughtful discussions concerning design and analysis
were provided by E. Marsh-Matthews, G. Wellborn and T. Wissing.
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